Passivation and remediation of Pb and Cr in contaminated soil by sewage sludge biochar tubule

Currently, numerous studies have carried out to research the effect of biochars remediation soil heavy metals (HMs) contaminated, but there have been fewer explorations of the effect of biochars tubule on soil HMs remediation. This work aimed to study the effect of passivation and remediation of lead (Pb) and chromium (Cr) contaminated soil after insert sewage sludge biochar (SSB) tubule. The results showed that the high risky fractions of Pb and Cr could be transformed into more stable fractions; also, Pb and Cr total contents are significantly decreased by SSB tubule. The mechanisms include adsorption, ion exchange, complexation, and precipitation which are concluded from the characteristic analysis. Detailly, the passivation of Pb and Cr is better when the moisture is 25% and 35%, respectively [Pb: exchangeable (F1), carbonate bound (F2) decreased by 25.1%, 16.8%, Fe-Mn oxides bound (F3) increased by 18.5%; Cr: F1 decreased by 73.0%, F2, F3, organic matter bound (F4) increased by 13.2%, 23.9%, 30.8%), respectively]. The remediation of Pb and Cr is better when the moisture is 25% and 35%, respectively, (Pb: decreased by 23.3%; Cr: decreased by 38.4%, respectively). The findings showed that the SSB tubule is effective when used for soil HMs contaminated.


Introduction
The soil environment has been a global concern due to its complexity and significance. However, with the rapid urbanization and industrialization, various contaminants [such as HMs, radioactive elements, and organics] are being introduced into the soil system through directly discharged to contain more than the allowable standard contaminant concentrations Rybak et al. 2018;Feng et al. 2018). HMs have ranked the first among all soil contaminant types due to their high toxicity, bioaccumulation, persistence, and mobility in soils Yang et al. 2020). These HMs (Pb, Cr, Cd, Hg, As, etc.) can lead to potential health risks to predators and humans through the accumulation of food chains. For instance, Pb is a commonly recognized carcinogen and led to neurotoxicity and stomach and lung lesions; Cr(VI) is a strong oxidant and also acts as carcinogenic and teratogenic characteristics, and both of them have been listed as priority monitoring and control pollutants . Excess of Pb and Cr in soil is mainly derived from intensifying anthropogenic activities including mining, sewage irrigation, and pesticide abuse , which could lead to a deterioration in various functions and stability of soil systems (Duan et al. 2018).
In the last several decades, various remediation technologies, including phytoremediation, excavation, landfilling, electrokinetic remediation, soil washing, and their blending, have been used to Pb-and Cr-contaminated soils for removing or reducing high toxic element amounts (Liu et al. 2018a;Sarwar et al. 2017;Trellu et al. 2016). However, most of these technologies due to their long remediation cycle, high energy consumption, low efficiency, and generating significant secondary environmental impacts (Gerhardt et al. 2009), are uns u i t a b l e f o r t h e m o s t a r e a t o d e c e l e r a t e t h e i r development. Accordingly, there is an urgent need for inexpensive, efficient, and stable amendment materials.
To date, various materials [resin (Chen et al. 2020a); activated carbon (Dong et al. 2016); peat (Lee et al. 2015)] are used to adsorb HMs. Unfortunately, the resin has a potential risk with secondary pollution; activated carbon is expensive and peat has a characteristically low surface. Biochar is a material with carbon-rich, porous, and high aromaticity which is obtained from the pyrolysis of biomass at relatively low temperature (<700°C) under the presence of limited oxygen Chen et al. 2020b). Biochar as an environmental sorbent has become one of the most attractive research hotspots due to having abundant raw materials, easy preparation, stable performance, etc. (Hung et al. 2020;Zhang et al. 2020a;Xiao et al. 2019). In addition, biochar can increase crop yield through improve soil fertility and remediate soil by immobilizing HMs (Azeem et al. 2021;Xi et al. 2020). Numerous researches have showed that biochar plays important role in decreasing the HMs (such as Cr, Cd, Pb, Cu, and Zn) total and unstable concentration (Puga et al. 2015;Gao et al. 2020;Liu et al. 2018b). Thus, biochar has great advantages as a green environmental sorbent in remediating HMscontaminated soil.
Biochar can be prepared from a wide variety of raw materials such as municipal sludge, livestock manure, and crop straw (Al-Wabel et al. 2018). In recent years, municipal sewage sludge has increased sharply with the increasing improvement of sewage treatment facilities (Zhou et al. 2020a); it is a by-product of the sewage treatment municipal and contains a variety of harmful substances (pathogens, refractory organics and toxic HMs, etc.), which is easy to cause secondary pollution (Zhou et al. 2020b). Pyrolysis of sewage sludge is promising since it enables to decrease the harmful substances and volume. Meanwhile, the solid carbonaceous residues after pyrolysis (SSB) may be used as the amendment of HMs in soil (Zhang et al. 2020b). At present, some researches have showed the excellent effect of sludge biochar on the remediation of HMs-contaminated soil. For example, Penido et al. (2019) observed a significant reduction of Cd, Pb, and Zn bioavailability in HMs-contaminated soil from a Zn-mining area by applying SSB. Fang et al. (2016) used the SSB to remediate soils that had been contaminated with cationic Pb(II), etc. and anionic Cr(VI), etc. respectively. However, there are still concerns regarding potential soil secondary contamination with toxic HMs due to SSB that were thoroughly mixed with soil and require further investigations to the assessment of long-term risks (Fang et al. 2016;Figueiredo et al. 2019). Thereby, it is requisite to obtain a piece of equipment which is able to separate biochar from the soil when remediation ended.
In this study, biochar derived from sewage sludge was placed in polymethyl methacrylate (PMMA) tubule with micropore and as an amendment inserted HMs-contaminated soils. To eliminate or decrease the risk of HMs secondary return to the soil, the biochar tubules were recycling when remediation ended. Herein, Pb and Cr were selected as representative of toxic HMs. The objectives of this study are to (1) investigate the effect of SSB on the fractions of Pb and Cr in soil and (2) determine the changes of the total content of Pb and Cr in soil with the remediation distance from the SSB tubule and time. The research results will provide some theoretical references for HMs-contaminated soil remediation and passivation by biochar in future research and application.

Biochar preparation
In this study, sewage sludge (SS) was collected from the dewatering workshop of the Xintian Wastewater Treatment Plant (XWTP) in Wanzhou District, Chongqing City, China, where an Orbal oxidation ditch wastewater treatment system is operated. The basic properties of sewage sludge were given in Table 1. The raw sewage sludge was dewatered through a vacuum pump (SHZ-DIII, Yuezong, China) and then was placed in an oven (DHG-9420A, Honghua, China) at a temperature of 103±2°C to constant weight. The dried samples were ground and passed through 20-mesh nylon sieves. The samples were thoroughly mixed and stored in plastic bags for further study. According to our previous study, SSB's pyrolysis at 500°C for 2 h had the best pore structures and adsorption capacity. In brief, approximate 50.0 g of sewage sludge samples loaded in the crucible and covered with fitting lids and alumina foils to achieve an oxygen-deficient environment were pyrolysed at 500°C for 2 h by the muffle furnace (SX-10-12, Yiheng, China). The obtained SSB was grounded to pass through 100-mesh nylon sieves, and dip in hydrochloric acid for 24 h to eliminate effects of surface and internal impurities, and improve the porous structure of SSB (Liu et al. 2018b), and then the SSB was cleaned with deionized water until reaching a neutral pH. Finally, SSB was dried and store in plastic bags.

Characterization of biochar
The yield was calculated from the ratio between the mass of SSB and the raw materials. Ash was measured as the residual remaining after heating to 800°C and maintaining for 4 h. The pH values were measured with a pH meter (FE28, Mettler Toledo, China) and the method from Wang et al. (2020). Levels of Pb and Cr were determined by inductively coupled plasma optical emission spectrometry (ICP-OES; Optima7000DV, Perkin Elmer, USA) after acid digestion according to Figueiredo et al. (2019). The surface structure of SSB was analyzed using a scanning electron microscopy (SEM; Supra55, Zeiss Germany). The surface functional groups of SSB were determined through Fourier transform infrared (FTIR) spectroscopy (Nicolet iS 10, Thermo Fisher Scientific, USA), the infrared spectra were obtained over the 4000-500 cm −1 . The crystal structure of the SSB was characterized by X-ray diffraction (XRD; D8 ADVANCE, Bruker, Germany).

Test soil collection and preparation
Surface soil (0-20 cm) was collected from Three Gorges Reservoir area located in Wanzhou District, Chongqing City, China (30°42′ 53″ N, 108°25′ 55″ E). The soil sample was thoroughly mixed and air-dried at room temperature; excess roots and gravels were removed and ground to pass through 20-mesh nylon sieves before initiating the experiments. The physiochemical properties of the soil were determined according to Chinese standard methods (Liu et al. 1996) and were given in Table 2.
To obtain a more significant remediation effect of SSB on Pb and Cr and reduce influences of other HMs in the soil, we replenish external Pb (Pb(NO 3 ) 2 ) or Cr (K 2 Cr 2 O 7 ) to the airdried soil to reach the first or second type of Risk Intervention Values for Soil Contamination of Development Land (RIVSCDL) of China, respectively. Meanwhile, the soil moisture content was controlled at 25% or 35% in each test group. Finally, all test soils were allowed to incubate for 7 days before the experiments. The type of soils required, and marking was given in Table 3.

Experimental design
The contaminated soils and SSB (SSB replaced with raw soil for control groups) were placed in PMMA column [H (height)=10 cm, D (diameter) = 30 cm] and PMMA tubule (H=10 cm, d=1.5 cm, many pores was full of the tubule wall, and it was surrounded with 120-mesh nylon sieves to separate SSB particles from soil), respectively. The tubule was located center of the column, and samples were sequential collected at equal distance and depth to measure the change of HMs concentration ( Fig. 1). To minimize the migration of Pb and Cr through the column sidewalls, PMMA columns were coated with a thin layer of paraffin wax and spread a layer of 1.5 cm gravel. Then, the soil was filled with layer upon layer and packed. The filling height up to 8.0 cm both soil and biochar and at a ratio of 1:600 (dry mass, biochar/soils). Finally, they were incubated in a lightproof, sealed, and room temperature sustained 35 days, and soil samples were collected every 5 days and 8 times in total. Approximate 0.5 g soil samples were collected to measure the total contents of Pb or Cr, and the needless soil was backfilled and compacted. After the last collection, SSB tubules were recycled, and soil in the PMMA column was thoroughly mixed and measured the Pb and Cr concentration of fractions. All of the soil samples were dried and grounded to pass through 100-mesh nylon sieves and stores. Each experiment was implemented in duplicate. The schematic diagram of the experimental design is shown in Fig. 1.

Total contents of Pb and Cr
In this study, total Pb and Cr were determined using a microwave digester (Speedwave Xpert, Berghof, Germany). Particularly, samples (0.2 g) were digested with a dup-acid mixture of HF (1 mL) and HNO 3 (8 mL) according to Khadhar et al. (2020) and the residual solution through membrane filters (0.22 μm) and diluted to 50 mL.

Chemical fractions of Pb and Cr
Pb and Cr potentially toxic were measured according to Tessier sequential extraction method (Tessier et al. 1979); this method divides HMs into five fractions due to successively less bioavailability and mobility effect (Chen et al. 2020c). In brief, extraction steps were as follows: (F1)

Passivation effect of SSB on Pb and Cr
HMs exist in the soil with diverse chemical fractions, and the bioavailability and potentially toxic mainly depend on their specific fraction. F1 presents the most unstable and high risky; F2, F3, and F4 are successive slightly; F5 is stable and safe (Chen et al. 2020c). Fraction distribution of Pb and Cr obtained from Tessier sequential extraction method is shown in Fig. 2 which reveal the fraction distribution difference of initially and finally. In brief, the results showed that CN had varied slightly of fraction distribution and SSB was significant. Compare SSB-F  (Fig. 2d). Munir et al. (2020) found that Cr and Pb be immobilized and removed by -OH and -COOH groups in biochar. Results showed that SSB could decrease Pb and Cr bioavailability and mobility through transforming the F1 and F2 to more stable forms and that is better when the moisture is 25% and 35%, respectively (Pb: F1, F2 decreased by 25.1%, 16.8%. Cr: F1 decreased by 73.0%, respectively). These results were compared with others. Liu et al. (2018b) added 5% (w/w) modified coconut shell biochar to multimetalcontaminated soil in Mianzhu, Sichuan, China, which decreased the acid-soluble Cd, Ni, and Zn concentration by 30.1%, 57.2%, and 12.7%, respectively. Munir et al. (2020) added 2% (w/w) bamboo biochar to Pb-and Cr-contaminated soil in Huainan, Anhui, China, which decreased their F1 concentration by 8.5% and 29%, respectively. Therefore, SSB tubules have a practical effect on soil passivation.

The remediation effect of SSB on Pb and Cr
The variation of Pb and Cr total contents in the soil with the incubation period and distance is shown in Fig. 3 (A line represents the variation of the Pb or Cr total content in this sampling point with time). These results showed that the total contents of Pb and Cr were significantly decreased. Presented different remediation results from different HMs and these results indicated SSB has a certain selectivity to HMs.
At end of the remediation, compared with initially, results revealed that CN had varied slightly of Pb and Cr concentration [concentration decreased by (−2.1)-8.0% and (−10.3)-4.0%]. On the contrary, the remediation effect was significantly of treatment groups. The Pb concentration decreased in the range of 8.5-21.8% with SSB-F and 14.6-23.3% with SSB-S, respectively (soil moisture content was 25%). Similarly, the Pb concentration decreased in the range of 13.0-17.3% and 8.2-16.2% when the soil moisture content was 35%. For Cr, the concentration decreased in the range of 9.5-22.2% and 5.0-15.5% when the soil moisture content was 25% and   Fig. 2 Fractions distribution of Pb (a-b) and Cr (c-d) in soil. F1, exchangeable; F2, carbonate bound; F3, Fe-Mn oxides bound; F4, organic matter bound; F5, residual. a and b: contaminated soil (Pb), at moisture 25% and 35%; c and d: contaminated soil (Cr), at moisture 25% and 35%; a and b: fractions distribution of Pb or Cr from initially and finally; CN, control. HMs-contaminated soil (concentration was controlled at the first type of RIVSCDL) which were remediated by raw soil. SSB-F and SSB-S: treatment groups. HMs-contaminated soil (concentration was controlled at the first and second type of RIVSCDL) which were remediated by SSB 25.1-38.4% and 24.8-36.1% when the soil moisture content was 35%. Maximum of Pb is located at the farthest and the closest point when moisture was 25% and 35%, respectively, and maximum of Cr located is shown in Fig. 3c-d. Results indicated that the Pb concentration maximum decreased by 23.3% when the soil moisture content was 25%, and the Cr concentration maximum decreased by 38.4% when the soil moisture content was 35%. This remediation effect was compared with others. Wang et al. (2020) added 10% (w/w) kitchen waste-based biochar to Ba-contaminated soil in landfill areas, Tibet, China, which decreased the concentration by 10.1%. Li et al. (2016) added biochar to Cd-contaminated soil and found that the total content decreased by 46.4%. These studies indicate that added biochars could remove HMs in soil.
In addition, for Pb-contaminated soil, these points (≤7.5 cm) presented initial decrease, subsequent increase, and final decrease with the remediation time, and others presented initial increase and final decrease. These phenomena could be attributed to SSB having a rich porous structure (Fig. 4b). The porous structure could provide more adsorption sites for HMs binding and provide accessible paths for HMs into the particle the biochar interior (Wang et al. 2018;Lin et al. 2020). Lin et al. (2020) indicated biochars adsorb HMs are divided into the surface monolayer sorption initially and the intraparticle diffusion sorption later. Initially, SSB adsorbs Pb ion at closer points, with the adsorption persistent; Pb ion could gradually migrate to SSB tubule from a broader area, which leads to Pb ion short-dated increase and final continuous decrease; Mitzia et al. (2020) consider that the cause of this phenomenon may attribute multiple interactions (such as soil matrix and water content), which can provoke an Eh-pH fluctuation. Crcontaminated soil presented a similar trend for all lines which are rapid initial decrease, subsequent increase (starting on the 5th day) and final decrease (starting on the 20th day) both presented an indistinctive effect at remediation distance.

Characteristics of SSB
The basic properties of SSB are given in Table 4. Results showed that yield and ash of SSB were higher compared with other biochar (biochars derived from rice straw, sawdust, and phragmites) which was mainly due to the SSB having high content of inorganic constituents (Pellera and Gidarakos 2015). Furthermore, SSB was alkaline which was due to the  surface of SSB including multitudinous alkaline aromatization groups, and they could immobilize HMs through increase the pH of the soil (Mitzia et al. 2020). Pellera and Gidarakos (2015) found that biochar leads to higher soil pH and ECE and causes metal precipitation, finally.

SEM analysis
The SEM was carried out to characterize the microstructure of SSB, and different magnifications are shown in Fig. 4. Results showed that the SSB had uneven size distribution and visible loose fold (Fig. 4a) and had a rich porous structure (Fig. 4b), which could be due to the breakdown of the volatile compounds at higher temperatures that lead to the energy (gas) to escape from sewage sludge (Shakya and Agarwal 2019). These form mesopores and micropores to obtain greater surface area and porous volume, and these porous structures could provide enough adsorption area for metal binding (Wang et al. 2018). Furthermore, visible inorganic ash particles were randomly distributed around of porous structure; Yuan et al. (2020) found that these particles would via ion exchange, complexation, and precipitation reactions to reduce metal ions.

FTIR spectra analysis
Figure 5 a shows the infrared spectrum of SSB. Five main absorption peaks centered at 3734, 2360, 1540, 1015, and 775 cm −1 were recorded. The peaks observed near 3734 cm −1 was attributed to the stretching and bending vibration of -OH (Lin et al. 2017), the peaks at 2360 cm −1 were corresponded to the stretching vibration of CO 2 (Lin et al. 2017), and the peaks at 1540 cm −1 was ascribed to the stretching vibration of C=O and C=C aromatic rings (Shin et al. 2020), and the peaks at 1015 cm −1 and 775 cm −1 were attributed to the stretching vibration of -OH and C-H of aromatic rings , respectively, indicating that SSB contained all kinds of functional groups. These functional groups play an important role to reduce Pb and Cr in soil that has been widely reported. For example, Zhao et al. (2021) found that O-containing functional groups were involved in Cr(VI) removal. Wang et al. (2018) found that -OH and C=O were involved in Pb(II) removal by the ion exchange reaction.

XRD analysis
Figure 5 b shows the mineralogical composition of SSB, illustrating highly crystalline structures due to revealing numerous sharp peaks . Results revealed C, SiO 2 , SiS 2 , and AlPO 4 as the predominant minerals in SSB. The characteristic peaks of carbon were detected corresponding to the presence of a number of aromatic carbon sheets, and this phenomenon was confirmed from the FTIR results. Also, the components of SiO 2 , SiS 2 , and AlPO 4 were responsible to immobilize Pb and Cr in soil. Li et al. (2018) found that SiO 2 could promote complexation and coordination with Cr(VI), the SiS 2 could induce Cr(VI) convert into Cr(III), and the AlPO 4 were converted to Pb 5 (PO 4 ) 3 OH which was precipitated (Zhao et al. 2018). Based on the previous discussions, the potential mechanism of removal and passivation of Pb and Cr is surmised. Generally, SSB could decrease Pb and Cr toxicity and mobility due to the formed precipitations under alkaline conditions (Khan et al. 2020); on the other hand, abundant pore structures of SSB provide more adsorption sites for metal complexing and precipitating (Wang et al. 2018). Furthermore, the FTIR spectrum revealed that functional groups (such as -OH, C=O, C=C, etc.) of SSB could bind Pb and Cr via ion exchange, complexation, and coordination (Yuan et al. 2020). XRD analysis indicated that SiO 2 and AlPO 4 could immobilize Pb and Cr via complexation and precipitation (Shakya and Agarwal 2019;Zhao et al. 2018). To sum up, these characters of SSB could reduce the bioavailability and total contents of Pb and Cr in soil.

Cost analysis of SSB
In this study, SSB was prepared in a muffle furnace with an oxygen-limited pyrolysis method. Specific costs of raw materials, reagents, and disposal were estimated, which provides a basis for large-scale production in the further. To date, more than 60 million tons of municipal sewage sludge was produced in China . The raw materials gathered were costless. Furthermore, the cost of electricity during the production procedures is 0.51 USD/kg (including pyrolysis and dry of biochar). Finally, the cost of chemicals (hydrochloric acid) during removing SSB impurities procedures is 0.71 USD/kg. The cost of SSB has a low level compared with anterior reports, and the price will lower when large-scale production (Cai et al. 2021). Therefore, the SSB has a great potential in HMs-contaminated soil remediation; meanwhile, there are signs to investigate further.

Conclusion
The fraction of unstable and high risky and total content of Pb and Cr could be significantly decreased when inserting a SSB tubule in the contaminated soil, which was achieved via adsorption, ion exchange, complexation, precipitation, etc. This study revealed that there is a better passivation effect of Pb and Cr when the soil moisture was 25% and 35%, respectively [Pb: (F1) decreased by 4.9%, (F2) decreased by 14.8%, (F3) increased by 22.5%; Cr: (F1) decreased by 60.2%, (F3) increased by 21.5%), maximally]. The remediation effect is better for Pb and Cr when the moisture content was 25% and 35% (the total content decreased by 23.3% and 38.4%, maximally) which is located at outermost and innermost, respectively. Furthermore, SSB exhibited higher efficacy in remediating and passivating Cr contamination than Pb.